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Contaminant Concentrations and Reproductive Rate of Lake Superior Bald Eagles, 1989–2001
J. Great Lakes Res. 31:227–235 Internat. Assoc. Great Lakes Res., 2005
Contaminant Concentrations and Reproductive Rate of Lake Superior Bald Eagles, 1989–2001 Cheryl R. Dykstra1,*, Michael W. Meyer2, Paul W. Rasmussen3, and D. Keith Warnke4,† 1Department
of Wildlife Ecology University of Wisconsin-Madison Madison, Wisconsin 53706 2Wisconsin
Department of Natural Resources 107 Sutliff Avenue Rhinelander, Wisconsin 54501
3Wisconsin
Department of Natural Resources 1359 Femrite Drive Monona, Wisconsin 537163736
4Department
of Fisheries and Wildlife Cooperative Research Unit University of Minnesota St. Paul, Minnesota 55108
ABSTRACT. We investigated the trend in contaminant concentrations in Lake Superior bald eagles (Haliaeetus leucocephalus) from 1989–2001, and examined the relationship of contaminant concentrations to eagle reproductive rate during that time. Concentrations of dichloro-diphenyl-dichloroethylene (DDE) and total polychlorinated biphenyls (PCBs) in nestling blood plasma samples decreased significantly from 1989–2001 (p = 0.007 for DDE, p = 0.004 for total PCBs). Mean contaminant concentrations in eaglet plasma, 21.7 µg/kg DDE (n=51) and 86.7 µg/kg total PCBs (n = 54), were near or below the estimated threshold levels for impairment of reproduction as determined in other studies. A preliminary assessment of polybrominated diphenyl ether (PBDE) concentrations indicated a mean of 7.9 µg/kg total PBDEs in Lake Superior eaglet plasma (n = 5). The number of occupied bald eagle nests along the Wisconsin shore of Lake Superior increased from 15 to 24 per year, between 1989 and 2001 (p < 0.001, r2 = 0.70, n = 13 years). Eagle reproductive rate did not increase or decrease significantly between 1989 and 2001 (p = 0.530, r2 = 0.037, n = 13 years, mean productivity = 0.96 young per occupied nest). The lack of correlation between reproductive rate and contaminant concentrations, as well as the comparison of contaminant concentrations to the estimated thresholds for impairment of reproduction, suggest that DDE and PCBs no longer limit the reproductive rate of the Lake Superior eagle population in Wisconsin. INDEX WORDS:
Bald eagle, DDE, PCB, PBDE, Lake Superior, reproductive rate.
INTRODUCTION Between 1947 and 1970, populations of bald eagles nesting in the continental United States declined due to reproductive failure caused mainly by
dichloro-diphenyl-dichloroethylene (DDE), a metabolite of the organochlorine insecticide dichloro-diphenyl-trichloroethane (DDT; Wiemeyer et al. 1972, Colborn 1991). After the use of DDT and other organochlorines was banned, the North American eagle population rebounded quickly throughout most of its range. However, bald eagle populations in a few regions, including the Great Lakes shorelines (Colborn 1991, Best et al. 1994) did not increase as rapidly.
*Corresponding author. E-mail: [email protected]. Present address: 7280 Susan Springs Drive, West Chester, OH 45069. †Present address: Wisconsin Department of Natural Resources, WM/6 Box 7921, Madison, WI, 53707.
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On the shorelines of Lake Superior in 1970, 100% of bald eagle reproductive efforts failed (Postupalsky 1971). The reproductive rate did not improve until 1977, when six young were fledged (Postupalsky 1978). Three addled eggs collected in 1970 contained 34, 65, and 71 mg/kg DDE (Postupalsky 1971), significantly more than the amounts associated with reproductive failure of eagles in other studies, > 3.6 mg/kg DDE or > 13 mg/kg total polychlorinated biphenyls (PCBs) in addled eggs (Wiemeyer et al. 1984, 1993), and 6 mg/kg DDE or 20 mg/kg total PCBs in addled eggs (calculated from plasma concentration thresholds for significant impairment of productivity, [Elliott and Harris 2001/2002]). Thus, organochlorine contaminants, particularly DDE, were implicated as the cause of the historic reproductive failure on Lake Superior. The principal source of organochlorine contaminants such as DDE and PCBs in Lake Superior was atmospheric deposition (Strachan and Eisenreich 1988, Arimoto 1989, Stevens and Neilson 1989). Since the 1970s, the concentrations of DDE and total PCBs in the biota of Lake Superior have decreased significantly (DeVault et al. 1986, Weseloh et al. 1994). Despite this decrease, contaminant concentrations in eagle eggs were still elevated in the mid-1980s and eagle reproduction along Lake Superior remained lower than that of neighboring eagles nesting in inland Wisconsin (Kozie and Anderson 1991), leading to the conclusion that organochlorine contaminants were still depressing productivity of Lake Superior eagles, at least in Apostle Islands National Lakeshore, in southern Lake Superior (Kozie 1986, Kozie and Anderson 1991). However, by the early 1990s, with contaminant concentrations in eagle eggs continuing to decrease (Dykstra et al. 1998), researchers concluded that the still-depressed reproductive rate of Wisconsin Lake Superior eagles was likely caused by low food availability rather than by organochlorine contaminants (Dykstra et al. 1998), although a contribution from contaminants, particularly DDE, could not be conclusively ruled out. Similarly, population modeling indicated that productivity of Apostle Island eagles was positively correlated with populations of prey fish species burbot (Lota lota) and longnose sucker (Catostomus catostomus) from 1983–1999 (Hoff et al. 2004). Organochlorine concentrations in some Great Lakes biota had stopped declining by the 1990s, having reached plateau concentrations which were still somewhat elevated in comparison to animals in other regions. Total PCBs and TCDD congeners in
herring gull (Larus argentatus) eggs from many Great Lakes locations were no longer decreasing by the mid-1980s to 1990s (Hebert et al. 1994, Stow 1995, Pekarik and Weseloh 1998). Similarly, DDE concentrations in lake trout (Salvelinus namaycush) from Lake Ontario declined rapidly in the 1970s but were relatively stable in the 1980s (Borgmann and Whittle 1991), and PCBs in seven fish species in Lake Michigan had reached stable concentrations by the late 1980s (Stow et al. 1995). However, the trends in contaminant concentrations in eagles might differ from those in fish and gulls in the same habitat because of the eagles’ position at the top of the food chain; it might be expected that contaminants would persist longer and at higher levels in a long-lived top predator such as eagles (Elliott and Harris 2001/2002). Additionally, in contaminant “hotspots” located in various regions of the Great Lakes, continued monitoring of contaminants with the potential to impact eagle reproduction, such as DDE, is prudent (Elliott and Harris 2001/2002). The purpose of this study was to investigate the trend in contaminant concentrations in Lake Superior eagles in the 1990s, and to examine the relationship of contaminant concentrations to eagle reproductive rate during that time. METHODS Contaminant Concentrations in Nestling Blood Between 1989 and 2001, blood samples were collected from 54 bald eagle nestlings or pairs of sibling nestlings at 25 nesting territories located within 8 km of the Lake Superior shore in Wisconsin (n = 24) or in Minnesota < 26 km from Wisconsin (n = 1). Ten of the 25 territories were located offshore on the Apostle Islands, 2 to 25 km from the mainland shore, and 15 were located along the mainland shore (Fig. 1). Nestlings were sexed by footpad length, and aged by the length of the eighth primary (Bortolotti 1984). Nestlings were 5–10 weeks old at the time of the blood collection. Syringes used were sterile plastic or glass previously washed with hexanes and acetone. Approximately 10 mL of blood was drawn from the brachial vein. Blood was transferred to heparinized vacutainers, stored on wet ice until the end of the day, and separated by centrifuging in the evening, or stored in a refrigerator for one day and then centrifuged less than 48 hours after collection. Plasma was drawn off, transferred to another vacutainer, and frozen upright at approximately –20°C. At the end of the field season, all frozen plasma
Contaminants and Reproduction in Lake Superior Eagles
FIG. 1. Map of the study area along the Wisconsin shore of Lake Superior. Triangle markers indicate bald eagle nest sites where nestling blood samples were collected, 1989–2001. samples were shipped on dry ice to Michigan State University (1989–1994) or Wisconsin State Laboratory of Hygiene (1998–2001). For samples collected 1989–1994 (n = 36), organochlorine pesticides and total PCB concentrations in the nestling plasma were determined by gas chromatography, with electron capture detection and were confirmed by mass spectrometry by the Michigan State University Aquatic Toxicology Laboratory (MSU-ATL hereafter). Methods have already been described (Bowerman 1993, Mora et al. 1993). Detection limits were 2.5 µg/kg for DDE and 5.0 µg/kg for total PCBs. Contaminant values falling below the detection limits were reported as one-half the detection limit (n = 3 DDE samples) and residue levels measured in the plasma of sibling nestlings in the same year were averaged to produce one value (n = 1). Samples collected in 1998–2001 (n = 18) were extracted at Wisconsin State Laboratory of Hygiene (WSLH hereafter) and analyzed on a gas chromatograph with electron-capture detection (Wisconsin State Laboratory of Hygiene 1996). Concentrations of p,p′- DDE and of 75 PCB congeners were measured; the sum of the PCB congener concentrations was considered to be the measure of Total PCBs. For five samples, concentrations of nine polybromi-
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nated diphenyl ether (PBDE) congeners were also measured at WSLOH; the sum of these was considered to be the Total PBDE concentration. The five samples analyzed for PBDEs were collected in 2000 and 2001 from one island nest (Michigan Island), three Wisconsin shoreline nests ranging from the eastern state line (Saxon Harbor) to the western state line (Superior), and one Minnesota shoreline nest (Sucker River). Detection limits were 0.15 µg/kg for DDE, 0.1–6 µg/kg for the PCB congeners, and 0.25–0.35 µg/kg for the PBDE congeners. No samples had non-detectable concentrations of DDE; individual PCB and PBDE congeners with concentrations below the detection limits were assigned a value of 0 for the purposes of summing congener concentrations. Laboratory results from the two facilities were comparable. Analysis of plasma PCBs and DDE at WSLH and MSU-ATL used similar extraction and clean-up procedures, GC columns, electron capture instrumentation, and Aroclor mixtures as standards. Spike recovery of p,p′-DDE from calf serum at WSLH was 77% for very low (0.5 µg/L) spike concentrations and 97% for spikes at 3.0 µg/L, similar to the range of recoveries (88–97%) reported by MSUATL for a mixture of chlorinated pesticides that included DDE. A larger number of p,p′-DDE spikes by WSLH (n = 27) yielded an average recovery of 92%, very close to the reported MSU-ATL average of 94% for pesticides. Spike recoveries of PCBs by the two labs were very similar, 90% for WSLH (average PCB congener recovery for two spikes), versus an average of 101% and range of 87% to 115% for total PCBs for MSU-ATL, and 87% average recovery of Aroclor 1254 for MSU-ATL. Reproductive Rate Reproductive rate was assessed by the Wisconsin Department of Natural Resources (WDNR) from 1989–2001 by inspecting nests from the air twice during the breeding season, once during incubation and again when nestlings were 4–7 weeks old. In the first aerial survey, the eagle pairs that were occupying territories and/or incubating eggs were counted, and in the second flight, the resulting nestlings were counted. For a regional summary, the total number of young produced was divided by the total number of territories that birds occupied. An occupied territory was defined as one where eggs had been laid, or two eagles were present on the territory, or the nest had been visibly repaired
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(even if no adults or only one non-incubating adult was observed; Postupalsky 1974). Statistical Analyses Concentrations of contaminants were log-transformed before analysis to correct for skewed distributions. Log-transformed values were used to calculate the geometric mean and 95% C.I. for residue concentrations in nestling plasma, and were used in analyses of time trends. We screened for potential outliers using the boxplot rule (Iglewicz and Hoaglin 1993), which identifies observations that are more than 1.5 times the interquartile range from either quartile as unusual observations. Multiple samples collected from eaglets at the same nests in different years cannot be assumed to be independent, because the eaglets likely have the same parents and because characteristics of the territory do not change much from year to year. To examine time trends in contaminant concentrations we used mixed effects models with a linear time trend (fixed effect) and random effects for nest sites. The random effects allow contaminant concentrations at each nest site to deviate consistently from the population mean but with a common trend over time for all sites. Such consistent differences among sites could be caused by differences in forage base, adult health or experience (if the site is used repeatedly by the same adults), or other factors not specified by the model. This model accounts for the lack of independence among observations from the same nest site by incorporating two error terms, one for variability among observations at the same site, and one for variability among sites. Note that linear trends in log concentration correspond to exponential trends in concentration. Computing was carried out using SAS PROC MIXED (Littell et al. 1996). Annual reproductive rates and the number of occupied nests per year were log-transformed to account for non-normality before analysis. Trends in the number of occupied nests and in reproductive rate were analyzed using linear regression. RESULTS Contaminant Concentrations in Nestling Blood The boxplot rule identified three DDE concentrations that were below the detection limit of 2.5 µg/kg as potential outliers. These three observations were all from 1991, they were substantially smaller than any other observation from these three nest sites (the next smallest observed DDE concentra-
FIG. 2. Concentrations of DDE in nestling blood plasma samples from bald eagle nests along the Wisconsin Lake Superior shore, 1989–2001. Y-axis is log-scale. Lines connect samples collected at the same nesting territory in different tion for these three nest sites was 14 µg/kg), and they were not associated with low PCB concentrations for the same samples. Thus these observations were excluded from further analyses. Geometric mean concentration of DDE in nestling blood during 1989–2001 was 21.7 µg/kg (95% C.I.= 18.0–26.1 µg/kg, n = 51 samples), but mean concentration for only those samples collected in 2000–2001 was lower, 13.4 µg/kg (95% C.I.= 10.2–17.7 µg/kg, n = 15 samples). Concentrations of DDE in nestling blood samples decreased significantly from 1989–2001, as evaluated by the mixed effects model (p = 0.007, slope = –0.0233, SE = 0.0079, df = 25, Fig. 2). The linear decrease corresponds to a decrease of 5.2% per year in DDE concentration. The variance among observations within nest sites was larger than that for consistent differences among nest sites (0.058 and 0.019, respectively). Concentrations of total PCBs in nestling blood samples averaged 86.7 µg/kg (95% C.I. = 70.4–107.0, n = 54 samples) for 1989–2001. Mean concentration for samples collected in 2000–2001 was lower, 51.5 µg/kg (95% C.I.= 40.0–66.3 µg/kg, n = 15). Total PCB concentrations decreased significantly from 1989–2001, as indicated by the mixed effects model (p = 0.004, slope = –0.0283, SE = 0.009, df = 28, Fig. 3). This linear decrease on the log scale corresponds to a decrease in PCB concentration of 6.3% per year. The variance among observations within nest sites was larger than that for
Contaminants and Reproduction in Lake Superior Eagles
FIG. 3. Concentrations of total PCBs in nestling blood plasma samples from bald eagle nests along the Wisconsin Lake Superior shore, 1989–2001. Y-axis is log-scale. Lines connect samples collected at the same nesting territory in different years. Value of the sample shown above the upper limit of graph was 1,154 µg/kg. consistent differences among nest sites (0.077 and 0.028, respectively). Nine different congeners of PBDE were measured in the plasma samples (Table 1). Of the nine congeners measured, BDE-47 and BDE-99 together accounted for the majority (36–49% and 23–33% respectively) of all PBDEs. BDE-100 was the next most abundant congener at 16–20%, while congeners BDE-153 & BDE-154 were detected in all samples but at ≤ 10% of the total. The concentra-
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FIG. 4. Number of occupied bald eagle nests along the Wisconsin shore of Lake Superior, 1989–2001. Line is the regression described in the text. tion of total PBDEs found in five Wisconsin Lake Superior bald eagle nesting plasma samples in 2000–2001 averaged 7.9 µg/kg (geometric mean; 95% C.I. = 6.0–10.4 µg/kg). Congeners BDE-28, BDE-66, BDE-85, and BDE-138 were not detected in any samples. The detection limits for BDE-66, BDE-85, and BDE-138 were < 0.3. However, interference from PCBs increased detection limits for BDE-28 to 0.6–1.0. Productivity The number of occupied nests along the Wisconsin shore of Lake Superior ranged from 15–24 per year between 1989 and 2001 and increased signifi-
TABLE 1. Congener specific PBDE concentrations (µg/kg) in Wisconsin Lake Superior bald eagle plasma samples (n = 5), 2000 - 2001. PBDE 2,4,4′ Tribromodiphenyl ether 2,2′,4,4′ Tetrabromodiphenyl ether 2,3′,4,4′ Tetrabromodiphenyl ether 2,2′,3,4,4′ Pentabromodiphenyl ether 2,2′,4,4′,5 Pentabromodiphenyl ether 2,2′,4,4′,6 Pentabromodiphenyl ether 2,2′,3,4,4′,5′ Hexabromodiphenyl ether 2,2′,4,4′,5,5′ Hexabromodiphenyl ether 2,2′,4,4′,5,6′ Hexabromodiphenyl ether Total PBDE (arithmetic mean) Total PBDE (geometric mean) 1. Arithmetic mean. 2. n.d. = not detected. 3. 95 % C.I.
Congener # #28 #47 #66 #85 #99 #100 #138 #153 #154 ∑ Congeners ∑ Congeners
Mean1 n.d.2 3.3 n.d. n.d. 2.4 1.5 n.d. 0.6 0.6 8.4 7.9
Range n.d. 2.5–5.1 n.d. n.d. 1.4–4.2 1.1–2.4 n.d. 0.6–1.0 0.4–0.9 6.1–13.6 6.0–10.43
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FIG. 5. Reproductive rate of bald eagles nesting along the Wisconsin shore of Lake Superior, 1989–2001. Reproductive rate measured as number of young produced per occupied nest, as defined in text. cantly during that time period (linear regression using log-transformed data, p < 0.001, r2 = 0.70, n = 13 years, Fig. 4). Reproductive rate ranged from 0.33 to 1.39 young per occupied nest (Fig. 5), with a mean production of 0.96 young per occupied nest. Reproductive rate did not increase or decrease significantly from 1989–2001 (p = 0.530, r2 = 0.037, n = 13 years). In 1994 and 1999, very low reproductive rates of 0.33 and 0.62 young per occupied territory, respectively, were recorded (Fig. 5). DISCUSSION Contaminant Concentrations Contaminant concentrations in eaglet blood samples generally reflect contaminant levels within the home territory (Olsson et al. 2000), as well as concentrations found in addled eggs from the same region (Elliott and Harris 2001/2002). Significant relationships have been derived between regional mean contaminant concentrations in eggs and plasma (Elliott and Norstrom 1998, Elliott and Harris 2001/2001). Using the relationship described in Elliott and Harris (2001/2002), we estimated that the mean 1989–2001 Lake Superior plasma concentrations of DDE and PCBs (21.7 µg/kg and 86.7 µg/kg, respectively) corresponded to egg values of approximately 3.9 mg/kg DDE and 10.7 mg/kg total PCBs. The mean concentrations in plasma at the end of our study period, 2000–2001, were equiva-
lent to about 2.8 mg/kg DDE and 7.1 mg/kg PCBs in addled eggs. The significance of contaminant levels for affecting eagle reproduction may be assessed by comparing samples to threshold values determined in other studies. In a recent comprehensive review, Elliott and Harris (2001/2002) estimated the thresholds for significant impairment of productivity to be 28 µg/kg DDE and 190 µg/kg total PCBs in nestling plasma, corresponding to 6 mg/kg DDE and 20 mg/kg total PCBs in addled eggs. In a previous study, contaminant concentrations > 3.6 mg/kg DDE or >13 mg/kg total PCBs in addled eggs were associated with bald eagle reproductive impairment (Wiemeyer et al. 1984, 1993). DDE concentrations in eagle nestling blood, and the calculated equivalent concentration in addled eggs, were near or below the threshold level for impairment of reproduction (depending on which egg threshold was used), suggesting that DDE might have depressed reproduction in the time from 1989 to 2001, but that the effect was likely slight. Additionally, because DDE levels in eaglets were decreasing from 1989 to 2001, and because by 2000–2001, the average DDE concentrations had dropped well below all threshold values, whether in plasma or calculated egg equivalents, we believe that DDE was likely not affecting eagle reproduction at that time. Mean total PCB concentrations in plasma and the calculated equivalent concentrations in addled eggs were below all thresholds for reproductive impairment, indicating that PCBs likely have had no effect on population productivity during the time period from 1989–2001. However, individual eaglets have recorded concentrations exceeding the thresholds, so possibly PCBs may have impaired reproduction at individual territories throughout the study period, without impacting the entire population. By 2000–2001, the mean contaminant levels were well below all suggested threshold values. DDE concentrations in plasma of Lake Superior eaglets in Wisconsin were similar to or lower than DDE levels elsewhere in the Great Lakes during the 1990s. Nestlings on the north shore of Lake Superior averaged 28 µg/kg DDE (1992–1994; Donaldson et al. 1999), and those on the north shore of Lake Erie 22 µg/kg DDE, (1990–1996; Donaldson et al. 1999). However, nestlings along the shore of Lake Michigan contained 48 µg/kg DDE (1987–2001; Dykstra et al. in press) and a single nestling sampled on Lake Huron had 63 µg/kg (Donaldson et al. 1999).
Contaminants and Reproduction in Lake Superior Eagles Total PCB concentrations in eaglets on the south shore of Lake Superior in the 1990s (86.7 µg/kg) were similar to those documented for north shore Lake Superior in the same time (94 µg/kg, 1992–1994; Donaldson et al. 1999). Along Lakes Erie and Michigan, concentrations were significantly higher (averaging 130 µg/kg and 233 µg/kg, respectively, Donaldson et al. 1999, Dykstra et al. in press). Polybrominated diphenyl ethers (PBDEs) are persistent and bioaccumulative compounds used as fire retardants in consumer goods, such as polyurethane foams, plastics, and textiles. PBDE concentrations have been documented in human tissue and biota from throughout the United States and Europe. This is the first measurement for Great Lakes eagles, and the data reported here should be considered preliminary due to the small sample size. PBDEs have been identified as a possible threat to aquatic biota, in part because of their prevalence and tendency to bioaccumulate (Luross et al. 2002), but relatively little is known about the toxicology of PBDEs. Concentrations of PBDEs in some Great Lakes biota have increased exponentially since 1981 (eggs of herring gulls, Norstrom et al. 2002). The PBDE concentrations in plasma of Lake Superior eaglets, 7.9 µg/kg, were lower than those measured in whole-fish samples of Lake Superior lake trout, 56 µg/kg (Luross et al. 2002), and lower than those measured in Lake Michigan salmon (80.1 µg/kg; Manchester-Neesvig et al. 2001). Lake Superior bald eaglet plasma PBDE concentrations were 10 times lower than those of double-crested cormorants (Phalacrocorax auritus) eggs collected in the nearby Lake Superior Apostle Island National Lakeshore (65–95 µg/kg fresh weight; S. Strom, WDNR pers. comm.), within the range of PBDE concentrations of fish collected in Wisconsin at sites remote from human/industrial activity (ND – 16 µg/kg fresh weight), but >10 times lower than PBDE concentrations measured in fish collected from rivers receiving treated wastewaters and land runoff in Wisconsin (50–482 µg/kg fresh weight; C. Schrank, WDNR pers. comm.) Reproductive Rate The mean reproductive rate of Wisconsin Lake Superior eagles in the 1990s, 0.96 young per occupied territory, is indicative of a healthy, expanding population (Buehler et al. 1991, Best et al. 1994, Bowman et al. 1995). Population expansion was evidenced by the significant increase in the number
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of occupied nesting territories in this region from 1989–2001. In fact, this population has grown substantially since 1974, when there were only two occupied territories along the Wisconsin Lake Superior shore (Kozie and Anderson 1991, Dykstra 1995, Dykstra et al. unpubl. data). Similar population growth and reproductive success have been documented in the eagle populations in other regions of the Lake Superior shore (Bowerman et al. 1998, Donaldson et al. 1999), indicating that the Wisconsin shoreline population is likely representative of the eagles of the entire lake. Reproductive rate did not increase through the 1990s, despite a decrease in both DDE and PCB concentrations, suggesting that there was no relationship between contaminant levels and productivity in this decade. Similarly, contaminant concentrations in eaglets were not correlated to productivity for individual territories along the Wisconsin Lake Superior shore in 1989–1993 (Dykstra et al. 1998). The lack of correlation between reproductive rate and contaminant concentrations, as well as the comparison of contaminant concentrations to the estimated thresholds for impairment of reproduction, suggest that DDE and PCBs no longer limit reproduction of the Lake Superior eagle population, at least in Wisconsin. However, reproduction at individual nests located in contaminant “hotspots” potentially may be affected by DDE or PCBs. Ecological Factors In the absence of anthropogenic factors such as chemical contaminants, variation in raptor reproductive rate (Fig. 5) usually results from ecological factors (Newton 1979), most typically weather and prey availability for bald eagles (reviewed in Elliott and Harris 2001/2002). For Lake Superior eagles, relatively high nestling mortality contributed to low reproductive rates in the mid-1980s (Kozie and Anderson 1991) and early 1990s (Dykstra et al. 1998); in the latter study, low prey delivery rates and low prey availability were shown to be the likely cause (Dykstra et al. 1998), although a contribution from contaminants, particularly DDE, could not be conclusively ruled out. Similarly, for 1983–1999, the productivity of Lake Superior eagles in the Apostle Islands was positively correlated with populations of two prey fish species there (Hoff et al. 2004). The effect of weather conditions may also be significant for Lake Superior eagles. For example, the very low eagle productivity in 1994 (Fig. 5) was
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preceded by an extremely cold winter (1993–1994) during which Lake Superior froze over completely. These conditions likely caused energetic stress in adult eagles by increasing their thermoregulatory energy demands and simultaneously reducing the availability of fish prey early in the breeding season. The cause of the unusually low productivity recorded in 1999 (Fig. 5) is unknown. CONCLUSIONS In summary, results from our study suggest that ecological factors, including food availability and possibly weather, are now more significant in limiting Lake Superior bald eagle reproductive rate than are the environmental contaminants DDE and PCBs. These factors must be investigated when attempting to understand eagle population dynamics for Lake Superior, and likely for the other Great Lakes as well. However, the presence of individual contaminant “hotspots” and the emergence of newly-discovered contaminant threats, such as PBDEs, indicate the need for continued monitoring and study. ACKNOWLEDGMENTS We thank Dave Evans for climbing to nests to retrieve the eaglets. Nest surveys in Wisconsin were conducted by Wisconsin Department of Natural Resources. Bill Bowerman, John Giesy, and colleagues, Michigan State University, analyzed some of the blood samples. Funding for this project was provided in part by the U.S. Fish and Wildlife Service, the Wisconsin Department of Natural Resources, and Wisconsin’s Great Lakes Protection Fund. REFERENCES Arimoto, R. 1989. Atmospheric deposition of chemical contaminants to the Great Lakes. J. Great Lakes Res. 15:339–356. Best, D.A., Bowerman, W.W., IV., Kubiak, T.J., Winterstein, S.R., Postupalsky, S., Shieldcastle, M.C., and Giesy, J.P., Jr. 1994. Reproductive impairment of Bald Eagles Haliaeetus leucocephalus along the Great Lakes shorelines of Michigan and Ohio. In Raptor Conservation Today, eds. B.-U. Meyburg and R.D. Chancellor , pp. 697–702, World Working Group on Birds of Prey. East Sussex, Great Britain: Pica Press. Borgmann, U., and Whittle, D.M. 1991. Contaminant concentrations trends in Lake Ontario lake trout (Salvelinus namaycush): 1977–1988. J. Great Lakes Res. 17:368–381. Bortolotti, G.R. 1984. Criteria for determining age and
sex of nestling bald eagles. J. Field Ornithol. 55:467–481. Bowerman, W.W., IV. 1993. Regulation of bald eagle (Haliaeetus leucocephalus) productivity in the Great Lakes basin: an ecological and toxicological approach. Ph.D. dissertation, Michigan State University. East Lansing, MI. ———, Best, D.A., Grubb, T.G., Zimmerman, G.M., and Giesy, J.P. 1998. Trends of contaminants and effects in bald eagles of the Great Lakes Basin. Environ. Monitor. Assess. 53:197–212. Bowman, T.D., Schempf, P.F., and Bernatowicz, J.A. 1995. Bald Eagle survival and population dynamics in Alaska after the Exxon Valdez oil spill. J. Wildl. Manage. 59:317–324. Buehler, D.A., Fraser, J.D., Seegar, J.K.D., Therres, G.D., and Byrd, M.A. 1991. Survival rates and population dynamics of Bald Eagles on Chesapeake Bay. J. Wildl. Manage. 55:608–613. Colborn, T. 1991. Epidemiology of Great Lakes bald eagles. J. Toxicol. Environ. Health 33:395–453. DeVault, D.S., Wilford, W.A., Hesselberg, R.J., Nortrupt, D.A., Rundberg, E.G.S., Alwan, A.K., and Bautista, C. 1986. Contaminant trends in lake trout (Salvelinus namaycush) from the upper Great Lakes. Arch. Environ. Contam. Toxicol. 15:349–356. Donaldson, G.M., Shutt, J.L., and Hunter, P. 1999. Organochlorine contamination in bald eagle eggs and nestlings from the Canadian Great Lakes. Arch. Environ. Contam. Toxicol. 36:70–80. Dykstra, C.J.R. 1995. Effects of contaminants, food availability and weather on the reproductive rate of Lake Superior Bald Eagles (Haliaeetus leucocephalus). Ph.D. dissertation, University of Wisconsin-Madison, Madison, WI. ———, Meyer, M.W., Warnke, D.K., Karasov, W.H., Andersen, D.E., Bowerman, W.W., IV., and Giesy, J.P. 1998. Low reproductive rates of Lake Superior bald eagles: low food delivery rates or environmental contaminants? J. Great Lakes Res. 24:32–44. ———, Meyer, M.W., Postupalsky, S., Stromborg, K.L., Warnke, D.K., and Eckstein, R.G. In Press. Bald eagles of Lake Michigan: ecology and contaminants. In The Lake Michigan ecosystem: ecology, health, and management, eds. T. Edsall and M. Munawar. Aquatic Ecosystem Health and Management Society, Burlington, Ontario, Canada. Elliott, J.E., and Harris, M.L. 2001/2002. An ecotoxicological assessment of chlorinated hydrocarbon effects on bald eagle populations. Reviews in Toxicology 4:1–60. ———, and Norstrom, R. 1998. Chlorinated hydrocarbon contaminants and productivity of bald eagle populations on the Pacific coast of Canada. Environ. Toxicol. Chem. 17:1142–1153. Hebert, C.E., Norstrom, R.J., Simon, M., Braune, B.M., Weseloh, D.V., and Macdonald, C.R. 1994. Temporal
Contaminants and Reproduction in Lake Superior Eagles trends and sources of PCDDs and PCDFs in the Great Lakes: herring gull egg monitoring, 1981–1991. Environ. Sci. Technol 28:1268–1277. Hoff, M.H., Meyer, M.W., Van Stappen, J., and Fratt, T.W. 2004. Relationships between bald eagle productivity and dynamics of fish populations and fisheries in the Wisconsin waters of Lake Superior, 1983–1999. J. Great Lakes Res. 30(Suppl. 1): 434–442. Iglewicz, B., and Hoaglin, D.C. 1993. How to detect and handle outliers. Milwaukee, WI: Quality Press and American Society for Quality Control. Kozie, K.D. 1986. Breeding and feeding ecology of bald eagles in the Apostle Islands National Lakeshore. M.S. thesis. University of Wisconsin-Stevens Point, Stevens Point, WI. ———, and Anderson, R.K. 1991. Productivity, diet, and environmental contaminants in bald eagles nesting near the Wisconsin shoreline of Lake Superior. Arch. Environ. Contam. Toxicol. 20:41–48. Littell, R.C., Milliken, G.A., Stroup, W.W., and Wolfinger, R.D. 1996. SAS System for Mixed Models. Cary, NC: SAS Institute Inc. Luross, J.M., Alaee, M., Sergeant, D.B., Cannon, C.M., Whittle D.M., Solomon, K.R., and Muir, D.C.G. 2002. Spatial distribution of polybrominated diphenyl ethers and polybrominated biphenyls in lake trout from the Laurentian Great Lakes. Chemosphere 46:665–672. Manchester-Neesvig, J.B., Valters, K., and Sonzogni, W.C. 2001. Comparison of polybrominated diphenyl ethers (PBDEs) and polychlorinated biphenyls (PCBs) in Lake Michigan salmonids. Environ. Sci. Technol. 35:1072–1077. Mora, M.A., Auman, H.J., Ludwig, J.P., Giesy, J.P., Verbrugge, D.A., and Ludwig, M.E. 1993. Polychlorinated biphenyls and chlorinated insecticides in plasma of Caspian terns: relationships with age, productivity and colony site tenacity in the Great Lakes. Arch. Environ. Contam. Toxicol. 24:320–331. Newton, I. 1979. Population Ecology of Raptors. Vermillion, South Dakota: Buteo Books. Norstrom, R.J., Simon, M., Moisey, J., Wakeford, B., and Weseloh, D.V. 2002. Geographical distribution (2000) and temporal trends (1981–2000) of brominated diphenyl ethers in Great Lakes herring gull eggs. Environ. Sci. Technol. 36:4783–4789. Olsson, A., Ceder, K., Bergman, A., and Helander, B. 2000. Nestling blood of the white-tailed sea eagle (Haliaeetus albicilla) as an indicator of territorial exposure to organohalogen compounds–an evaluation. Environ. Sci. Technol. 34:2733–2740. Pekarik, C., and Weseloh, D.V. 1998. Organochlorine Contaminants in Herring Gull Eggs from the Great Lakes, 1974–1995: Change Point Regression Analysis and Short-Term Regression. Environ. Monitor. Assess. 53:77–115.
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Postupalsky, S. 1971. Toxic chemicals and declining bald eagles and cormorants in Ontario. Unpublished report to the Canadian Wildlife Service, No. 20. ———. 1974. Raptor reproductive success: some problems with methods, criteria, and terminology. In Management of Raptors, eds. F.N. Hamerstrom, Jr., B.E. Harrell, and R.R. Ohlendorff, pp. 21–31. Proc. Conf. Raptor Conserv. Tech., Raptor Res. Report No. 2. ———. 1978. The bald eagles return. Natural History 87(7):62–63. Stevens, R.J.J., and Neilson, M.A. 1989. Inter- and intralake distributions of trace organic contaminants in surface waters of the Great Lakes. J. Great Lakes Res. 15:377–393. Stow, C.A. 1995. Great Lakes herring gull egg PCB concentrations indicate approximate steady-state conditions. Environ. Sci. Technol. 29:2893–2897. ———, Carpenter, S.R., Eby, L.A., Amrhein, J.F., and Hesselberg, R.J. 1995. Evidence that PCBs are approaching stable concentrations in Lake Michigan fishes. Ecological Applications 5:248–260. Strachan, W., and Eisenreich, S. 1988. Mass balancing of toxic chemicals in the Great Lakes: the role of atmospheric deposition. Appendix 1 of the Workshop on Great Lakes Atmospheric Deposition. International Joint Commission. Weseloh, D.V., Ewing, P.J., Struger, J., Mineau, P., and Norstrom, R.J. 1994. Geographical distribution of organochlorine contaminants and reproductive parameters in herring gulls on Lake Superior in 1983. Environ. Monitor. Assess. 29:229–251. Wiemeyer, S.N., Mulhern, B.M., Ligas, F.J., Hensel, R.J., Mathisen, J.E., Robards, F.C., and Postupalsky, S. 1972. Residues of organochlorine pesticides, polychlorinated biphenyls, and mercury in bald eagle eggs and changes in shell thickness, 1969 and 1970. Pestic. Monitor. J. 6:50–55. ——— , Lamont, T.G., Bunck, C.M., Sindelar, C.R., Gramlich, F.J., Fraser, J.D., and Byrd, M.A. 1984. Organochlorine pesticide, polychlorinated biphenyls, and mercury residues in bald eagle eggs— 1969–1979—and their relationship to shell thinning and reproduction. Arch. Environ. Contam. Toxicol.13:529–549. ———, Bunck, C.M., and Stafford, C.J. 1993. Environmental contaminants in bald eagle eggs—1980–84— and further interpretations of relationships to productivity and shell thickness. Arch. Environ. Contam. Toxicol. 24:213-227. Wisconsin State Laboratory of Hygiene 1996. Methods of Organic Analysis Manual—January 1, 1996. Wisconsin State Laboratory of Hygiene, Madison, WI. Submitted: 18 November 2004 Accepted: 16 March 2005 Editorial : Martin A. Stapanian
Contaminant Concentrations and Reproductive Rate of Lake Superior Bald Eagles, 1989–2001 Cheryl R. Dykstra1,*, Michael W. Meyer2, Paul W. Rasmussen3, and D. Keith Warnke4,† 1Department
of Wildlife Ecology University of Wisconsin-Madison Madison, Wisconsin 53706 2Wisconsin
Department of Natural Resources 107 Sutliff Avenue Rhinelander, Wisconsin 54501
3Wisconsin
Department of Natural Resources 1359 Femrite Drive Monona, Wisconsin 537163736
4Department
of Fisheries and Wildlife Cooperative Research Unit University of Minnesota St. Paul, Minnesota 55108
ABSTRACT. We investigated the trend in contaminant concentrations in Lake Superior bald eagles (Haliaeetus leucocephalus) from 1989–2001, and examined the relationship of contaminant concentrations to eagle reproductive rate during that time. Concentrations of dichloro-diphenyl-dichloroethylene (DDE) and total polychlorinated biphenyls (PCBs) in nestling blood plasma samples decreased significantly from 1989–2001 (p = 0.007 for DDE, p = 0.004 for total PCBs). Mean contaminant concentrations in eaglet plasma, 21.7 µg/kg DDE (n=51) and 86.7 µg/kg total PCBs (n = 54), were near or below the estimated threshold levels for impairment of reproduction as determined in other studies. A preliminary assessment of polybrominated diphenyl ether (PBDE) concentrations indicated a mean of 7.9 µg/kg total PBDEs in Lake Superior eaglet plasma (n = 5). The number of occupied bald eagle nests along the Wisconsin shore of Lake Superior increased from 15 to 24 per year, between 1989 and 2001 (p < 0.001, r2 = 0.70, n = 13 years). Eagle reproductive rate did not increase or decrease significantly between 1989 and 2001 (p = 0.530, r2 = 0.037, n = 13 years, mean productivity = 0.96 young per occupied nest). The lack of correlation between reproductive rate and contaminant concentrations, as well as the comparison of contaminant concentrations to the estimated thresholds for impairment of reproduction, suggest that DDE and PCBs no longer limit the reproductive rate of the Lake Superior eagle population in Wisconsin. INDEX WORDS:
Bald eagle, DDE, PCB, PBDE, Lake Superior, reproductive rate.
INTRODUCTION Between 1947 and 1970, populations of bald eagles nesting in the continental United States declined due to reproductive failure caused mainly by
dichloro-diphenyl-dichloroethylene (DDE), a metabolite of the organochlorine insecticide dichloro-diphenyl-trichloroethane (DDT; Wiemeyer et al. 1972, Colborn 1991). After the use of DDT and other organochlorines was banned, the North American eagle population rebounded quickly throughout most of its range. However, bald eagle populations in a few regions, including the Great Lakes shorelines (Colborn 1991, Best et al. 1994) did not increase as rapidly.
*Corresponding author. E-mail: [email protected]. Present address: 7280 Susan Springs Drive, West Chester, OH 45069. †Present address: Wisconsin Department of Natural Resources, WM/6 Box 7921, Madison, WI, 53707.
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On the shorelines of Lake Superior in 1970, 100% of bald eagle reproductive efforts failed (Postupalsky 1971). The reproductive rate did not improve until 1977, when six young were fledged (Postupalsky 1978). Three addled eggs collected in 1970 contained 34, 65, and 71 mg/kg DDE (Postupalsky 1971), significantly more than the amounts associated with reproductive failure of eagles in other studies, > 3.6 mg/kg DDE or > 13 mg/kg total polychlorinated biphenyls (PCBs) in addled eggs (Wiemeyer et al. 1984, 1993), and 6 mg/kg DDE or 20 mg/kg total PCBs in addled eggs (calculated from plasma concentration thresholds for significant impairment of productivity, [Elliott and Harris 2001/2002]). Thus, organochlorine contaminants, particularly DDE, were implicated as the cause of the historic reproductive failure on Lake Superior. The principal source of organochlorine contaminants such as DDE and PCBs in Lake Superior was atmospheric deposition (Strachan and Eisenreich 1988, Arimoto 1989, Stevens and Neilson 1989). Since the 1970s, the concentrations of DDE and total PCBs in the biota of Lake Superior have decreased significantly (DeVault et al. 1986, Weseloh et al. 1994). Despite this decrease, contaminant concentrations in eagle eggs were still elevated in the mid-1980s and eagle reproduction along Lake Superior remained lower than that of neighboring eagles nesting in inland Wisconsin (Kozie and Anderson 1991), leading to the conclusion that organochlorine contaminants were still depressing productivity of Lake Superior eagles, at least in Apostle Islands National Lakeshore, in southern Lake Superior (Kozie 1986, Kozie and Anderson 1991). However, by the early 1990s, with contaminant concentrations in eagle eggs continuing to decrease (Dykstra et al. 1998), researchers concluded that the still-depressed reproductive rate of Wisconsin Lake Superior eagles was likely caused by low food availability rather than by organochlorine contaminants (Dykstra et al. 1998), although a contribution from contaminants, particularly DDE, could not be conclusively ruled out. Similarly, population modeling indicated that productivity of Apostle Island eagles was positively correlated with populations of prey fish species burbot (Lota lota) and longnose sucker (Catostomus catostomus) from 1983–1999 (Hoff et al. 2004). Organochlorine concentrations in some Great Lakes biota had stopped declining by the 1990s, having reached plateau concentrations which were still somewhat elevated in comparison to animals in other regions. Total PCBs and TCDD congeners in
herring gull (Larus argentatus) eggs from many Great Lakes locations were no longer decreasing by the mid-1980s to 1990s (Hebert et al. 1994, Stow 1995, Pekarik and Weseloh 1998). Similarly, DDE concentrations in lake trout (Salvelinus namaycush) from Lake Ontario declined rapidly in the 1970s but were relatively stable in the 1980s (Borgmann and Whittle 1991), and PCBs in seven fish species in Lake Michigan had reached stable concentrations by the late 1980s (Stow et al. 1995). However, the trends in contaminant concentrations in eagles might differ from those in fish and gulls in the same habitat because of the eagles’ position at the top of the food chain; it might be expected that contaminants would persist longer and at higher levels in a long-lived top predator such as eagles (Elliott and Harris 2001/2002). Additionally, in contaminant “hotspots” located in various regions of the Great Lakes, continued monitoring of contaminants with the potential to impact eagle reproduction, such as DDE, is prudent (Elliott and Harris 2001/2002). The purpose of this study was to investigate the trend in contaminant concentrations in Lake Superior eagles in the 1990s, and to examine the relationship of contaminant concentrations to eagle reproductive rate during that time. METHODS Contaminant Concentrations in Nestling Blood Between 1989 and 2001, blood samples were collected from 54 bald eagle nestlings or pairs of sibling nestlings at 25 nesting territories located within 8 km of the Lake Superior shore in Wisconsin (n = 24) or in Minnesota < 26 km from Wisconsin (n = 1). Ten of the 25 territories were located offshore on the Apostle Islands, 2 to 25 km from the mainland shore, and 15 were located along the mainland shore (Fig. 1). Nestlings were sexed by footpad length, and aged by the length of the eighth primary (Bortolotti 1984). Nestlings were 5–10 weeks old at the time of the blood collection. Syringes used were sterile plastic or glass previously washed with hexanes and acetone. Approximately 10 mL of blood was drawn from the brachial vein. Blood was transferred to heparinized vacutainers, stored on wet ice until the end of the day, and separated by centrifuging in the evening, or stored in a refrigerator for one day and then centrifuged less than 48 hours after collection. Plasma was drawn off, transferred to another vacutainer, and frozen upright at approximately –20°C. At the end of the field season, all frozen plasma
Contaminants and Reproduction in Lake Superior Eagles
FIG. 1. Map of the study area along the Wisconsin shore of Lake Superior. Triangle markers indicate bald eagle nest sites where nestling blood samples were collected, 1989–2001. samples were shipped on dry ice to Michigan State University (1989–1994) or Wisconsin State Laboratory of Hygiene (1998–2001). For samples collected 1989–1994 (n = 36), organochlorine pesticides and total PCB concentrations in the nestling plasma were determined by gas chromatography, with electron capture detection and were confirmed by mass spectrometry by the Michigan State University Aquatic Toxicology Laboratory (MSU-ATL hereafter). Methods have already been described (Bowerman 1993, Mora et al. 1993). Detection limits were 2.5 µg/kg for DDE and 5.0 µg/kg for total PCBs. Contaminant values falling below the detection limits were reported as one-half the detection limit (n = 3 DDE samples) and residue levels measured in the plasma of sibling nestlings in the same year were averaged to produce one value (n = 1). Samples collected in 1998–2001 (n = 18) were extracted at Wisconsin State Laboratory of Hygiene (WSLH hereafter) and analyzed on a gas chromatograph with electron-capture detection (Wisconsin State Laboratory of Hygiene 1996). Concentrations of p,p′- DDE and of 75 PCB congeners were measured; the sum of the PCB congener concentrations was considered to be the measure of Total PCBs. For five samples, concentrations of nine polybromi-
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nated diphenyl ether (PBDE) congeners were also measured at WSLOH; the sum of these was considered to be the Total PBDE concentration. The five samples analyzed for PBDEs were collected in 2000 and 2001 from one island nest (Michigan Island), three Wisconsin shoreline nests ranging from the eastern state line (Saxon Harbor) to the western state line (Superior), and one Minnesota shoreline nest (Sucker River). Detection limits were 0.15 µg/kg for DDE, 0.1–6 µg/kg for the PCB congeners, and 0.25–0.35 µg/kg for the PBDE congeners. No samples had non-detectable concentrations of DDE; individual PCB and PBDE congeners with concentrations below the detection limits were assigned a value of 0 for the purposes of summing congener concentrations. Laboratory results from the two facilities were comparable. Analysis of plasma PCBs and DDE at WSLH and MSU-ATL used similar extraction and clean-up procedures, GC columns, electron capture instrumentation, and Aroclor mixtures as standards. Spike recovery of p,p′-DDE from calf serum at WSLH was 77% for very low (0.5 µg/L) spike concentrations and 97% for spikes at 3.0 µg/L, similar to the range of recoveries (88–97%) reported by MSUATL for a mixture of chlorinated pesticides that included DDE. A larger number of p,p′-DDE spikes by WSLH (n = 27) yielded an average recovery of 92%, very close to the reported MSU-ATL average of 94% for pesticides. Spike recoveries of PCBs by the two labs were very similar, 90% for WSLH (average PCB congener recovery for two spikes), versus an average of 101% and range of 87% to 115% for total PCBs for MSU-ATL, and 87% average recovery of Aroclor 1254 for MSU-ATL. Reproductive Rate Reproductive rate was assessed by the Wisconsin Department of Natural Resources (WDNR) from 1989–2001 by inspecting nests from the air twice during the breeding season, once during incubation and again when nestlings were 4–7 weeks old. In the first aerial survey, the eagle pairs that were occupying territories and/or incubating eggs were counted, and in the second flight, the resulting nestlings were counted. For a regional summary, the total number of young produced was divided by the total number of territories that birds occupied. An occupied territory was defined as one where eggs had been laid, or two eagles were present on the territory, or the nest had been visibly repaired
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(even if no adults or only one non-incubating adult was observed; Postupalsky 1974). Statistical Analyses Concentrations of contaminants were log-transformed before analysis to correct for skewed distributions. Log-transformed values were used to calculate the geometric mean and 95% C.I. for residue concentrations in nestling plasma, and were used in analyses of time trends. We screened for potential outliers using the boxplot rule (Iglewicz and Hoaglin 1993), which identifies observations that are more than 1.5 times the interquartile range from either quartile as unusual observations. Multiple samples collected from eaglets at the same nests in different years cannot be assumed to be independent, because the eaglets likely have the same parents and because characteristics of the territory do not change much from year to year. To examine time trends in contaminant concentrations we used mixed effects models with a linear time trend (fixed effect) and random effects for nest sites. The random effects allow contaminant concentrations at each nest site to deviate consistently from the population mean but with a common trend over time for all sites. Such consistent differences among sites could be caused by differences in forage base, adult health or experience (if the site is used repeatedly by the same adults), or other factors not specified by the model. This model accounts for the lack of independence among observations from the same nest site by incorporating two error terms, one for variability among observations at the same site, and one for variability among sites. Note that linear trends in log concentration correspond to exponential trends in concentration. Computing was carried out using SAS PROC MIXED (Littell et al. 1996). Annual reproductive rates and the number of occupied nests per year were log-transformed to account for non-normality before analysis. Trends in the number of occupied nests and in reproductive rate were analyzed using linear regression. RESULTS Contaminant Concentrations in Nestling Blood The boxplot rule identified three DDE concentrations that were below the detection limit of 2.5 µg/kg as potential outliers. These three observations were all from 1991, they were substantially smaller than any other observation from these three nest sites (the next smallest observed DDE concentra-
FIG. 2. Concentrations of DDE in nestling blood plasma samples from bald eagle nests along the Wisconsin Lake Superior shore, 1989–2001. Y-axis is log-scale. Lines connect samples collected at the same nesting territory in different tion for these three nest sites was 14 µg/kg), and they were not associated with low PCB concentrations for the same samples. Thus these observations were excluded from further analyses. Geometric mean concentration of DDE in nestling blood during 1989–2001 was 21.7 µg/kg (95% C.I.= 18.0–26.1 µg/kg, n = 51 samples), but mean concentration for only those samples collected in 2000–2001 was lower, 13.4 µg/kg (95% C.I.= 10.2–17.7 µg/kg, n = 15 samples). Concentrations of DDE in nestling blood samples decreased significantly from 1989–2001, as evaluated by the mixed effects model (p = 0.007, slope = –0.0233, SE = 0.0079, df = 25, Fig. 2). The linear decrease corresponds to a decrease of 5.2% per year in DDE concentration. The variance among observations within nest sites was larger than that for consistent differences among nest sites (0.058 and 0.019, respectively). Concentrations of total PCBs in nestling blood samples averaged 86.7 µg/kg (95% C.I. = 70.4–107.0, n = 54 samples) for 1989–2001. Mean concentration for samples collected in 2000–2001 was lower, 51.5 µg/kg (95% C.I.= 40.0–66.3 µg/kg, n = 15). Total PCB concentrations decreased significantly from 1989–2001, as indicated by the mixed effects model (p = 0.004, slope = –0.0283, SE = 0.009, df = 28, Fig. 3). This linear decrease on the log scale corresponds to a decrease in PCB concentration of 6.3% per year. The variance among observations within nest sites was larger than that for
Contaminants and Reproduction in Lake Superior Eagles
FIG. 3. Concentrations of total PCBs in nestling blood plasma samples from bald eagle nests along the Wisconsin Lake Superior shore, 1989–2001. Y-axis is log-scale. Lines connect samples collected at the same nesting territory in different years. Value of the sample shown above the upper limit of graph was 1,154 µg/kg. consistent differences among nest sites (0.077 and 0.028, respectively). Nine different congeners of PBDE were measured in the plasma samples (Table 1). Of the nine congeners measured, BDE-47 and BDE-99 together accounted for the majority (36–49% and 23–33% respectively) of all PBDEs. BDE-100 was the next most abundant congener at 16–20%, while congeners BDE-153 & BDE-154 were detected in all samples but at ≤ 10% of the total. The concentra-
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FIG. 4. Number of occupied bald eagle nests along the Wisconsin shore of Lake Superior, 1989–2001. Line is the regression described in the text. tion of total PBDEs found in five Wisconsin Lake Superior bald eagle nesting plasma samples in 2000–2001 averaged 7.9 µg/kg (geometric mean; 95% C.I. = 6.0–10.4 µg/kg). Congeners BDE-28, BDE-66, BDE-85, and BDE-138 were not detected in any samples. The detection limits for BDE-66, BDE-85, and BDE-138 were < 0.3. However, interference from PCBs increased detection limits for BDE-28 to 0.6–1.0. Productivity The number of occupied nests along the Wisconsin shore of Lake Superior ranged from 15–24 per year between 1989 and 2001 and increased signifi-
TABLE 1. Congener specific PBDE concentrations (µg/kg) in Wisconsin Lake Superior bald eagle plasma samples (n = 5), 2000 - 2001. PBDE 2,4,4′ Tribromodiphenyl ether 2,2′,4,4′ Tetrabromodiphenyl ether 2,3′,4,4′ Tetrabromodiphenyl ether 2,2′,3,4,4′ Pentabromodiphenyl ether 2,2′,4,4′,5 Pentabromodiphenyl ether 2,2′,4,4′,6 Pentabromodiphenyl ether 2,2′,3,4,4′,5′ Hexabromodiphenyl ether 2,2′,4,4′,5,5′ Hexabromodiphenyl ether 2,2′,4,4′,5,6′ Hexabromodiphenyl ether Total PBDE (arithmetic mean) Total PBDE (geometric mean) 1. Arithmetic mean. 2. n.d. = not detected. 3. 95 % C.I.
Congener # #28 #47 #66 #85 #99 #100 #138 #153 #154 ∑ Congeners ∑ Congeners
Mean1 n.d.2 3.3 n.d. n.d. 2.4 1.5 n.d. 0.6 0.6 8.4 7.9
Range n.d. 2.5–5.1 n.d. n.d. 1.4–4.2 1.1–2.4 n.d. 0.6–1.0 0.4–0.9 6.1–13.6 6.0–10.43
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FIG. 5. Reproductive rate of bald eagles nesting along the Wisconsin shore of Lake Superior, 1989–2001. Reproductive rate measured as number of young produced per occupied nest, as defined in text. cantly during that time period (linear regression using log-transformed data, p < 0.001, r2 = 0.70, n = 13 years, Fig. 4). Reproductive rate ranged from 0.33 to 1.39 young per occupied nest (Fig. 5), with a mean production of 0.96 young per occupied nest. Reproductive rate did not increase or decrease significantly from 1989–2001 (p = 0.530, r2 = 0.037, n = 13 years). In 1994 and 1999, very low reproductive rates of 0.33 and 0.62 young per occupied territory, respectively, were recorded (Fig. 5). DISCUSSION Contaminant Concentrations Contaminant concentrations in eaglet blood samples generally reflect contaminant levels within the home territory (Olsson et al. 2000), as well as concentrations found in addled eggs from the same region (Elliott and Harris 2001/2002). Significant relationships have been derived between regional mean contaminant concentrations in eggs and plasma (Elliott and Norstrom 1998, Elliott and Harris 2001/2001). Using the relationship described in Elliott and Harris (2001/2002), we estimated that the mean 1989–2001 Lake Superior plasma concentrations of DDE and PCBs (21.7 µg/kg and 86.7 µg/kg, respectively) corresponded to egg values of approximately 3.9 mg/kg DDE and 10.7 mg/kg total PCBs. The mean concentrations in plasma at the end of our study period, 2000–2001, were equiva-
lent to about 2.8 mg/kg DDE and 7.1 mg/kg PCBs in addled eggs. The significance of contaminant levels for affecting eagle reproduction may be assessed by comparing samples to threshold values determined in other studies. In a recent comprehensive review, Elliott and Harris (2001/2002) estimated the thresholds for significant impairment of productivity to be 28 µg/kg DDE and 190 µg/kg total PCBs in nestling plasma, corresponding to 6 mg/kg DDE and 20 mg/kg total PCBs in addled eggs. In a previous study, contaminant concentrations > 3.6 mg/kg DDE or >13 mg/kg total PCBs in addled eggs were associated with bald eagle reproductive impairment (Wiemeyer et al. 1984, 1993). DDE concentrations in eagle nestling blood, and the calculated equivalent concentration in addled eggs, were near or below the threshold level for impairment of reproduction (depending on which egg threshold was used), suggesting that DDE might have depressed reproduction in the time from 1989 to 2001, but that the effect was likely slight. Additionally, because DDE levels in eaglets were decreasing from 1989 to 2001, and because by 2000–2001, the average DDE concentrations had dropped well below all threshold values, whether in plasma or calculated egg equivalents, we believe that DDE was likely not affecting eagle reproduction at that time. Mean total PCB concentrations in plasma and the calculated equivalent concentrations in addled eggs were below all thresholds for reproductive impairment, indicating that PCBs likely have had no effect on population productivity during the time period from 1989–2001. However, individual eaglets have recorded concentrations exceeding the thresholds, so possibly PCBs may have impaired reproduction at individual territories throughout the study period, without impacting the entire population. By 2000–2001, the mean contaminant levels were well below all suggested threshold values. DDE concentrations in plasma of Lake Superior eaglets in Wisconsin were similar to or lower than DDE levels elsewhere in the Great Lakes during the 1990s. Nestlings on the north shore of Lake Superior averaged 28 µg/kg DDE (1992–1994; Donaldson et al. 1999), and those on the north shore of Lake Erie 22 µg/kg DDE, (1990–1996; Donaldson et al. 1999). However, nestlings along the shore of Lake Michigan contained 48 µg/kg DDE (1987–2001; Dykstra et al. in press) and a single nestling sampled on Lake Huron had 63 µg/kg (Donaldson et al. 1999).
Contaminants and Reproduction in Lake Superior Eagles Total PCB concentrations in eaglets on the south shore of Lake Superior in the 1990s (86.7 µg/kg) were similar to those documented for north shore Lake Superior in the same time (94 µg/kg, 1992–1994; Donaldson et al. 1999). Along Lakes Erie and Michigan, concentrations were significantly higher (averaging 130 µg/kg and 233 µg/kg, respectively, Donaldson et al. 1999, Dykstra et al. in press). Polybrominated diphenyl ethers (PBDEs) are persistent and bioaccumulative compounds used as fire retardants in consumer goods, such as polyurethane foams, plastics, and textiles. PBDE concentrations have been documented in human tissue and biota from throughout the United States and Europe. This is the first measurement for Great Lakes eagles, and the data reported here should be considered preliminary due to the small sample size. PBDEs have been identified as a possible threat to aquatic biota, in part because of their prevalence and tendency to bioaccumulate (Luross et al. 2002), but relatively little is known about the toxicology of PBDEs. Concentrations of PBDEs in some Great Lakes biota have increased exponentially since 1981 (eggs of herring gulls, Norstrom et al. 2002). The PBDE concentrations in plasma of Lake Superior eaglets, 7.9 µg/kg, were lower than those measured in whole-fish samples of Lake Superior lake trout, 56 µg/kg (Luross et al. 2002), and lower than those measured in Lake Michigan salmon (80.1 µg/kg; Manchester-Neesvig et al. 2001). Lake Superior bald eaglet plasma PBDE concentrations were 10 times lower than those of double-crested cormorants (Phalacrocorax auritus) eggs collected in the nearby Lake Superior Apostle Island National Lakeshore (65–95 µg/kg fresh weight; S. Strom, WDNR pers. comm.), within the range of PBDE concentrations of fish collected in Wisconsin at sites remote from human/industrial activity (ND – 16 µg/kg fresh weight), but >10 times lower than PBDE concentrations measured in fish collected from rivers receiving treated wastewaters and land runoff in Wisconsin (50–482 µg/kg fresh weight; C. Schrank, WDNR pers. comm.) Reproductive Rate The mean reproductive rate of Wisconsin Lake Superior eagles in the 1990s, 0.96 young per occupied territory, is indicative of a healthy, expanding population (Buehler et al. 1991, Best et al. 1994, Bowman et al. 1995). Population expansion was evidenced by the significant increase in the number
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of occupied nesting territories in this region from 1989–2001. In fact, this population has grown substantially since 1974, when there were only two occupied territories along the Wisconsin Lake Superior shore (Kozie and Anderson 1991, Dykstra 1995, Dykstra et al. unpubl. data). Similar population growth and reproductive success have been documented in the eagle populations in other regions of the Lake Superior shore (Bowerman et al. 1998, Donaldson et al. 1999), indicating that the Wisconsin shoreline population is likely representative of the eagles of the entire lake. Reproductive rate did not increase through the 1990s, despite a decrease in both DDE and PCB concentrations, suggesting that there was no relationship between contaminant levels and productivity in this decade. Similarly, contaminant concentrations in eaglets were not correlated to productivity for individual territories along the Wisconsin Lake Superior shore in 1989–1993 (Dykstra et al. 1998). The lack of correlation between reproductive rate and contaminant concentrations, as well as the comparison of contaminant concentrations to the estimated thresholds for impairment of reproduction, suggest that DDE and PCBs no longer limit reproduction of the Lake Superior eagle population, at least in Wisconsin. However, reproduction at individual nests located in contaminant “hotspots” potentially may be affected by DDE or PCBs. Ecological Factors In the absence of anthropogenic factors such as chemical contaminants, variation in raptor reproductive rate (Fig. 5) usually results from ecological factors (Newton 1979), most typically weather and prey availability for bald eagles (reviewed in Elliott and Harris 2001/2002). For Lake Superior eagles, relatively high nestling mortality contributed to low reproductive rates in the mid-1980s (Kozie and Anderson 1991) and early 1990s (Dykstra et al. 1998); in the latter study, low prey delivery rates and low prey availability were shown to be the likely cause (Dykstra et al. 1998), although a contribution from contaminants, particularly DDE, could not be conclusively ruled out. Similarly, for 1983–1999, the productivity of Lake Superior eagles in the Apostle Islands was positively correlated with populations of two prey fish species there (Hoff et al. 2004). The effect of weather conditions may also be significant for Lake Superior eagles. For example, the very low eagle productivity in 1994 (Fig. 5) was
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preceded by an extremely cold winter (1993–1994) during which Lake Superior froze over completely. These conditions likely caused energetic stress in adult eagles by increasing their thermoregulatory energy demands and simultaneously reducing the availability of fish prey early in the breeding season. The cause of the unusually low productivity recorded in 1999 (Fig. 5) is unknown. CONCLUSIONS In summary, results from our study suggest that ecological factors, including food availability and possibly weather, are now more significant in limiting Lake Superior bald eagle reproductive rate than are the environmental contaminants DDE and PCBs. These factors must be investigated when attempting to understand eagle population dynamics for Lake Superior, and likely for the other Great Lakes as well. However, the presence of individual contaminant “hotspots” and the emergence of newly-discovered contaminant threats, such as PBDEs, indicate the need for continued monitoring and study. ACKNOWLEDGMENTS We thank Dave Evans for climbing to nests to retrieve the eaglets. Nest surveys in Wisconsin were conducted by Wisconsin Department of Natural Resources. Bill Bowerman, John Giesy, and colleagues, Michigan State University, analyzed some of the blood samples. Funding for this project was provided in part by the U.S. Fish and Wildlife Service, the Wisconsin Department of Natural Resources, and Wisconsin’s Great Lakes Protection Fund. REFERENCES Arimoto, R. 1989. Atmospheric deposition of chemical contaminants to the Great Lakes. J. Great Lakes Res. 15:339–356. Best, D.A., Bowerman, W.W., IV., Kubiak, T.J., Winterstein, S.R., Postupalsky, S., Shieldcastle, M.C., and Giesy, J.P., Jr. 1994. Reproductive impairment of Bald Eagles Haliaeetus leucocephalus along the Great Lakes shorelines of Michigan and Ohio. In Raptor Conservation Today, eds. B.-U. Meyburg and R.D. Chancellor , pp. 697–702, World Working Group on Birds of Prey. East Sussex, Great Britain: Pica Press. Borgmann, U., and Whittle, D.M. 1991. Contaminant concentrations trends in Lake Ontario lake trout (Salvelinus namaycush): 1977–1988. J. Great Lakes Res. 17:368–381. Bortolotti, G.R. 1984. Criteria for determining age and
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Postupalsky, S. 1971. Toxic chemicals and declining bald eagles and cormorants in Ontario. Unpublished report to the Canadian Wildlife Service, No. 20. ———. 1974. Raptor reproductive success: some problems with methods, criteria, and terminology. In Management of Raptors, eds. F.N. Hamerstrom, Jr., B.E. Harrell, and R.R. Ohlendorff, pp. 21–31. Proc. Conf. Raptor Conserv. Tech., Raptor Res. Report No. 2. ———. 1978. The bald eagles return. Natural History 87(7):62–63. Stevens, R.J.J., and Neilson, M.A. 1989. Inter- and intralake distributions of trace organic contaminants in surface waters of the Great Lakes. J. Great Lakes Res. 15:377–393. Stow, C.A. 1995. Great Lakes herring gull egg PCB concentrations indicate approximate steady-state conditions. Environ. Sci. Technol. 29:2893–2897. ———, Carpenter, S.R., Eby, L.A., Amrhein, J.F., and Hesselberg, R.J. 1995. Evidence that PCBs are approaching stable concentrations in Lake Michigan fishes. Ecological Applications 5:248–260. Strachan, W., and Eisenreich, S. 1988. Mass balancing of toxic chemicals in the Great Lakes: the role of atmospheric deposition. Appendix 1 of the Workshop on Great Lakes Atmospheric Deposition. International Joint Commission. Weseloh, D.V., Ewing, P.J., Struger, J., Mineau, P., and Norstrom, R.J. 1994. Geographical distribution of organochlorine contaminants and reproductive parameters in herring gulls on Lake Superior in 1983. Environ. Monitor. Assess. 29:229–251. Wiemeyer, S.N., Mulhern, B.M., Ligas, F.J., Hensel, R.J., Mathisen, J.E., Robards, F.C., and Postupalsky, S. 1972. Residues of organochlorine pesticides, polychlorinated biphenyls, and mercury in bald eagle eggs and changes in shell thickness, 1969 and 1970. Pestic. Monitor. J. 6:50–55. ——— , Lamont, T.G., Bunck, C.M., Sindelar, C.R., Gramlich, F.J., Fraser, J.D., and Byrd, M.A. 1984. Organochlorine pesticide, polychlorinated biphenyls, and mercury residues in bald eagle eggs— 1969–1979—and their relationship to shell thinning and reproduction. Arch. Environ. Contam. Toxicol.13:529–549. ———, Bunck, C.M., and Stafford, C.J. 1993. Environmental contaminants in bald eagle eggs—1980–84— and further interpretations of relationships to productivity and shell thickness. Arch. Environ. Contam. Toxicol. 24:213-227. Wisconsin State Laboratory of Hygiene 1996. Methods of Organic Analysis Manual—January 1, 1996. Wisconsin State Laboratory of Hygiene, Madison, WI. Submitted: 18 November 2004 Accepted: 16 March 2005 Editorial : Martin A. Stapanian